Considerations for reducing food system energy demand while scaling up urban agriculture

Proponents suggest a number of energy-related benefits are realized through the reintroduction of food production within cities (Howe and Wheeler 1999, Garnett 1997, Smit and Nasr 1992, Kulak et al 2013). Studies most commonly highlight savings in transportation energy, reduced storage requirements at the wholesale/resale level, and energy inputs of food waste/loss along the supply chain, but also include additional biomass provision from silviculture (i.e. to offset energy imports; Smit and Nasr 1992), easier exploitation of resource use (Zhang et al 2013), and lower resource-intensity of production (Kulak et al 2013). Meanwhile, peri-urban agriculture can preserve higher-yielding prime agricultural land (Krannich 2006, Francis et al 2012), which has the potential to provide less resource-intensive production. Looking at more sophisticated integrated operations (vertical farms, integrated greenhouses), exploited waste streams (CO₂, heat, macronutrients) could offset energy requirements that are required for providing these inputs in conventional operations (Despommier 2013, Zhang et al 2013). Additionally, if the distributed nature of UA can be supported by a similarly distributed energy infrastructure system, food/agriculture waste can be digested locally to generate biogas for heat or electricity production, further decreasing the energy footprint of UA. Energy-related benefits associated with the various structures/locations of UA have also been described in table 1 (excluding transportation).

Energy demand in irrigation systems are a noteworthy component of scaled-up UA that must be considered in order to avoid inadvertently increasing demand relative to conventional open-field systems. Irrigation systems in an open-field agricultural setting are relatively low-energy when compared with potable urban water systems that could be used in UA; in one study open-field irrigation energy demand is estimated at 0.63 MJ m⁻³ water (Esengun et al 2007, used in the absence of a similar US case study). However, in a UA system, potable water may be used for irrigation and generally requires substantially more energy for treatment, with the Electric Power Research Institute (2002) suggesting an estimate of 1.3 MJ m⁻³ and 1.7 MJ m⁻³ for public utilities using surface and groundwater, respectively (including distribution), for a hypothetical 10 million gallon per day treatment plant. Meanwhile, Racoviceanu et al (2007) estimate energy demand at 2.3−2.5 MJ m⁻³ treated water used in the City of Toronto’s water treatment. The Racoviceanu et al (2007) study considers a surface water source, and includes chemical fabrication/transportation, treatment, and onsite pumping, though most of total energy intensity (~70%) is attributable to untreated and treated water pumping. Data on Massachusetts’ 2007 energy demand for water treatment and distribution suggests an average value of 1.4 MJ m⁻³ (US Environmental Protection Agency 2008), whereas California’s 2005 report on the energy-water relationship provides estimates of 1.4 MJ m⁻³ and 9.7 MJ m⁻³ for Northern and Southern California, respectively (range attributable to differences in energy required for conveyance from source to treatment facilities; Klein et al 2005). This latter California report also suggests that when desalination options are employed in water treatment, an additional 9.3−15.7 MJ m⁻³ and 3.7−9.3 MJ m⁻³ are required for seawater and brackish groundwater, respectively. It is worth noting that depth of groundwater source, pumping requirements for surface/groundwater, and on-farm treatment will influence the energy demand and could bring this figure closer in line with that from water utilities.

The types of secondary energy used can also vary for different types of irrigation, influencing both cost, overall energy efficiency, and GHG emissions. For example, Ontario, Canada’s field crop irrigation is typically powered by diesel systems, while greenhouse irrigation is generally powered by electricity (Carol 2010). Diesel has an emissions intensity of 74 kg CO₂e GJ⁻¹, while electricity grid GHG intensity in Ontario was 14 kg CO₂e GJ⁻¹ in 2014 (IPCC 2006, chapter 3). For comparison, US electricity emissions intensities ranged from 1–266 kg CO₂e GJ⁻¹ in 2012 (US EPA 2015).

Water use can be mitigated through the use of more water-efficient growing systems (such as hydroponic systems), though these can result in increased energy demand in pumping and lighting, and associated GHG emissions. For example, hydroponic¹³ systems have been shown to have lower water demand than soil-based production, in addition to avoiding the need for a solid growing medium and the associated energy inputs of its provision (Albaho et al 2008). However, Barbosa et al (2015) have modeled energy and water demand for hydroponic and conventional production systems for lettuce; while water demand was reduced by 92% (250–20 l kg⁻¹ y⁻¹), energy demand increased by 8100% (1100 to 90 000 kJ kg⁻¹ y⁻¹), due primarily to heating and cooling loads (74 000 kJ kg⁻¹ y⁻¹), artificial lighting (15 000 kJ kg⁻¹ y⁻¹) and circulating pumps (640 kJ kg⁻¹ y⁻¹).

Focusing on energy, Shiina et al (2011) study hydroponic urban ‘plant factories’ (temperature controlled, artificial lighting and humidity controlled) in Japan, and show that the energy intensity of the production resulted in estimated greenhouse emissions of 6.4 kg CO₂e kg⁻¹ lettuce, despite the operation’s high yields. Continuing to use GHG emissions as a proxy for energy demand, this compares with estimates of 0.2 and 0.9 kg CO₂e kg⁻¹ for lettuce from Michigan hoop houses and California open-field lettuce production (Plawecki et al 2014), and ranges between 0.24−2.62 kg CO₂e kg⁻¹ for lettuce from European open field and hothouse production (Hospido et al 2009). Meanwhile, Goldstein et al (2016a) compared cumulative energy demand of rooftop hydroponic greenhouse tomatoes and ‘conventional’ production and find the former to be roughly ten times as energy intensive, with important implications for carbon footprint. However, switching energy source from the Massachusetts electricity grid to hydroelectric or solar PV makes rooftop hydroponic greenhouse production less carbon intensive than conventional production.

These demonstrate that are potential for trade-offs when addressing environmental footprints through UA if focusing on a single performance metric (i.e. water alone). Though, as hydroponic growing systems can be used in controlled, protected, and open-field growing systems and with a wide selection of hydroponic technology options available, variation can be expected in the yields and energy demand of hydroponic operations; this introduces uncertainty in applying these figures to specific contexts, but underscores the need for careful consideration in designing for energy and water demand reduction.

Urban agricultural systems provide an application for rainwater collection, as well as black/greywater¹⁴, all of which could reduce wastewater volumes and stormwater runoff, and potentially improve surface water quality and decrease net energy use as a result (i.e. due to the avoidance of UA irrigation with potable water and downstream wastewater treatment). As examples, wastewater treatment in California and Massachusetts is estimated to require, on average, 1.7 and 2.4 MJ m⁻³, respectively (US Environmental Protection Agency 2008, Klein et al 2005). This has the potential to be reduced if conveyance and treatment requirements are avoided through application of wastewater in UA. Further, if stormwater can be diverted from treatment plants to UA in jurisdictions using combined sewer systems, energy demand, as well as pollutants to receiving bodies, could be reduced. In an extreme case, substantial diversion of rainwater for UA from lakes and rivers that ordinarily receive it could contribute to local/regional ecosystem decline or surface water quality issues (Goldstein et al 2016a). Finally, depending on how UA is managed, runoff from open field urban farms could result in increased nutrient loads being passed down to receiving bodies or downstream wastewater treatment plants (Pataki et al 2011). Upscaling UA could result in this being an additional source of non-point pollution for consideration by city managers/planners.

The use of packaging materials can also potentially be avoided in UA operations, in instances of production for personal consumption or within shorter distribution chains such as when food is sold directly by the producer (Garnett 1999). For example, the climate impacts of the embodied energy of polyethylene terephthalate clamshells and polystyrene trays that are often used in tomato packaging (again, using carbon as a proxy for energy use) were estimated to be 25% and 100% greater, respectively, per unit mass of tomato when compared to loose packaging (US Environmental Protection Agency 2010). Still, the authors noted that modified atmosphere packaging using plastics have been shown to increase shelf life by two or three times, which may reduce waste and, consequently, GHGs associated with tomato production and disposal. This waste reduction could then offset the embodied energy needed for the packaging material that provides this added shelf life.

The use of packaging does not need to be an all or nothing proposition; employing some packaging for various meal components can result in a net energy savings (relative to ‘typical’ packaging configurations) when accounting for avoided waste and marginal energy requirements; semi-prepared meals examined by Hanssen et al (2017) were slightly more energy efficient when compared with those prepared from scratch. It is generally important to recognize the embodied energy of the food products and packaging materials being considered; higher embodied energy food products (cheese, beef, bread) more easily justifying the additional energy inputs associated with packaging than unprocessed fruits and vegetables (Williams and Wikstrom 2011). Similarly, the application of plastic films and containers may be more easily justified when compared with more energy-intensive materials such as steel, aluminum, or glass.

While UA and other forms of localization are often intuitively thought to reduce life cycle energy demand, the reality is more complicated (Webb et al 2013). Supply chains crossing a variety of artificial jurisdictional boundaries may in fact be more direct than those created by constraining agriculture within a region/state, depending on the product, consumption point, and regional characteristics (Nicholson et al 2015). Broad-scale localization of agriculture has the potential to increase transportation energy, as well as associated GHG emissions, relative to the conventional supply chain if definitions of local and implications for modified supply networks, including transport modes, are not carefully considered. Indeed, a commonly cited reason to pursue UA is to reduce energy-related impacts associated with transportation. Estimates of transportation’s contribution to the food system’s energy demand and GHG emissions have been estimated at approximately 10% or less (Weber and Matthews 2008, USDA ERS 2010, Garnett 2011).

Numerous studies from the literature (Coley et al 2009, Edwards-Jones et al 2008, Pirog et al 2001) have challenged the common assumption that ‘localizing’ food production results in reduced transport energy use and GHG emissions, and effects on distribution networks need to be evaluated on a case basis to justify such a claim. For instance, transport-related impacts for cheese shipped 20 000 km from New Zealand to consumers in England by boat were dominated by road-freight and consumer automobile use, highlighting the limitations of singular focus on transport distance (Basset-Mens et al 2007). The GHG implications of external energy inputs to support year-round urban food production and their ability to overwhelm gains achieved through reduced distribution distances must be considered in the context of upscaling of urban food production.

The predominance of dark (low-albedo) surfaces in cities results in the absorption of solar radiation and elevated temperatures in and around urban areas, raising the demand for cooling energy (the urban heat island effect; Oke 1973). Urban agriculture could play a role in attenuating this phenomenon, by increasing surface albedo and the cooling effect of plant evapotranspiration (Ackerman et al 2014). Vegetation situated on buildings has been shown to reduce individual building cooling demands in Toronto, Canada, Madrid, Spain and La Rochelle, France (Bass and Baskaran 2001, Saiz et al 2006, Jaffal et al 2012). Ackermann and colleagues estimated that scaling up UA in New York City could reduce the local urban heat island by 22%−44% (~1 °C), mitigating energy demands for cooling (Ackerman 2012). The importance of this ancillary benefit of UA could become more important with the increasing frequency and severity of heat waves under likely climate change scenarios (Jansson 2013).

Judging the pressures production systems have on resource demands requires reflection on a number of contextual factors. For example, local climate/geography may reduce the need for energy-intensive inputs (i.e. mild climate, plentiful surface/rain water). As well, existing infrastructure (green and grey) may or may not provide access to necessary inputs (nutrients, water, energy, labor, and growing media). This reflection may also include questions such as whether there is an abundance of low-grade heat that is accessible for exploitation and is the supplier (i.e. a local utility) amenable to supporting its exploitation, or perhaps if there is an existing agreement to supply nutrients from wastewater to peri-urban agriculture or further afield. Additionally, an abundance of uncontaminated vacant land or a low population density may make open-field or protected systems the most plausible approach. Further considerations with respect to publically-owned land might be whether these local green spaces are compatible with UA integration, when safety, waste collection, accessibility, and public demand are taken into account. Finally, Pelletier et al (2011) suggest that scale of production systems may also play a role in energy efficiency; though scale in itself is not an indicator of energy efficient production, smaller operations have been observed to have lower energy intensities in the examples of tomatoes and swine. It is clear that further research is needed to parse out the roles that scale, climate, existing infrastructure, waste resource availability can have on the overall energy picture of UA operations. Moreover, an assessment of the local context is necessary before promoting any particular UA approach, along with the accompanying resource demands these systems require in a given context.

Food waste has the potential to be converted to a useful energy resource in the form of biogas, with many cities already collecting source-separated organics for processing in local anaerobic digesters (Uçkun Kiran et al 2014, Sanscartier et al 2012, Mohareb et al 2011, Bernstad and la Cour Jansen 2011). Following the potential for circular resource use suggested by Metson et al (2012), the proximity of increased urban food waste from both production as well as further down the food supply chain could provide a greater feedstock for co-located urban anaerobic digestion (AD) systems. In addition, digestate produced from these facilities could find local end-uses in UA operations, facilitating a circular material flow. Governments are currently promoting UA to reduce the carbon footprint of cities (Arup and C40 Cities 2014). Keeping this objective in mind, it is important to consider how food waste (a major component of GHG emissions from landfills; US EPA 2017) can be better utilized within a more cyclical UA system.

Using food waste for energy generation through AD provides an opportunity for distributed energy generation while decreasing the impact of food waste on downstream systems (landfills, wastewater treatment plants). Levis and Barlaz (2011) assessed the environmental performance of food waste disposal in nine common waste management systems and found that AD performed best with respect to GHG emissions, NO_x, SO₂ and net energy demand. Further, considering the proximity to potential end users, the use of biogas from AD facilities for both heat and electricity production could become more economically attractive in an urban context, especially with local UA consumers of waste CO₂ (from biogas production) and AD digestate. It is estimated that the US cities produce 130 Mt of food waste annually¹⁶. Using estimates of 184 kWh of electricity and 810 MJ heat Mg⁻¹ of wet waste (from Møller et al 2009), this quantity of food waste has the potential to provide electricity for 7.2 million Nissan Leaf all-electric vehicles¹⁷ and the equivalent heating demand for over 1.5 million Michigan homes¹⁸, respectively.

Cities are currently operating AD facilities that are providing energy to the broader community. Barcelona is treating 192 000 t yr⁻¹ of its organic fraction of municipal solid waste (OFMSW) through AD, having a positive energy balance of around 2.2 MJ produced/MJ consumed at the facility from pre-treatments and digester pumping/stirring (Romero-Guiza et al 2014). Additionally, anaerobic co-digestion with sewage sludge could enhance biogas production and deals with the seasonality that food waste from UA can present (Fonoll et al 2015, Shrestha et al 2017). Policy interventions will likely be necessary to encourage broader investment in AD (Binkley et al 2013). For example, in the north of Italy, 26 000–28 000 of OFMSW are treated each year in AD plant; while the facility has obtained a positive cash flow of €2.5 million yr⁻¹, an incentive for the use/generation of renewable energy was needed to enable this to occur (Riva et al 2014).

Beyond energy production, AD offers additional benefits. Situating anaerobic digesters near UA operations could facilitate the reuse of digestate (such as in Garfí et al 2011), saving on fertilizer requirements and reducing transportation costs for waste diversion. The coupling of AD with pyrolysis has the potential to produce biochar, which could be used to improve soil fertility (Monlau et al 2016). Excess heat from AD or pyrolysis can also be applied to the digester to or to district heating systems and can be used to heat houses or aquaculture operations.

The barriers associated with the reintroduction of livestock into relatively dense areas are formidable; these include local regulations, public health concerns, and logistic difficulties of feed provision (Food and Agriculture Organization 2001, Butler 2011). If surmounted, these operations, as well as primary and secondary food processing industries (e.g. breweries, ethanol production, harvest-related waste from agricultural operations) can provide substantial feedstocks for AD.

Finally, in cases where AD is impractical, UA provides a local end user for composted residues. Hence, onsite compost facilities could be a component of future UA operations. This would reduce GHG emissions from waste that would have been disposed of in a landfill, and avoids the need for transportation of waste to a location offsite. According to the US EPA WARM model¹⁹, composting food waste and avoiding its addition to landfill results in a net reduction of 0.96 Mg CO₂e per Mg of food waste.

Both solid and liquid streams of wastewater are an underutilized resource, with their current perception as a municipal liability requiring resource-intensive treatment and disposal. It has been estimated that approximately 2% of the total US electricity use is for municipal wastewater treatment (Electric Power Research Institute 2002). The aeration step of treatment, which promotes biodegradation of pollutants, accounts for approximately 50% of this energy use (Curtis 2010, Mamais et al 2015). This approach also results in the release of GHG emissions to the atmosphere; in 2000, US wastewater treatment resulted in ~33.3 Mt CO₂e from energy use and sludge degradation (Center for Sustainable Systems 2014). A system that diverts wastewater from treatment, reduces the level of treatment, or eliminates the need for aeration (through diversion from receiving water bodies to UA) could help reduce these emissions.

Wastewater reuse could be a practical source of water and nutrients in UA. Previous studies have noted heavy metal and pathogen contamination of wastewater-irrigated produce (Amoah et al 2007, Khan et al 2008), underscoring the need to ensure regulatory requirements for irrigation water quality are met (World Health Organization 2006). If cities/neighborhoods were to reorient their wastewater treatment goals from a focus on disposal to one of reuse, the treatment reduction could result in substantial energy savings—directly at the point of treatment, as well as upstream from crop production. For example, crops grown using water and nutrients recovered from wastewater could offset the embodied energy demand of crops that are grown elsewhere using more energy-intensive irrigation water and inorganic fertilizers. Anaerobic membrane bioreactors are one technology that has been proposed to accomplish these goals (Smith et al 2012, 2014), recovering energy, generating an effluent rich in nutrients and low in suspended solids and organics, and eliminating energy requirements related to aerobic treatment (Smith et al 2014). Regardless of the technology used, further research is necessary to evaluate the removal potential of trace contaminants and viral pathogens prior to reuse for UA (Smith et al 2012, McCurry et al 2014). By taking an industrial ecology approach, residential waste streams and industrial waste streams that are relatively benign and with a low pathogen load (e.g. brewery waste) could be used in subsurface irrigation of UA crops, avoiding conventional treatment and reclaiming nutrients for food production.

Finally, a further industrial ecological approach would see conventional infrastructure systems integrated with agriculture to increase productivity. Many sources of waste heat and CO₂ exist within the urban boundary, from residences to industrial operations to electrical utilities. Where natural gas is employed in these applications, greenhouse operations can utilize the relatively clean exhausted low-grade energy as a heat source, as well as CO₂ for crop fertilization (Kimball 1983, Mortensen 1987). If greenhouses and households could be integrated, there is a potential efficiency gain in the combined system over its discrete components, including through the provision of CO₂ for crop fertilization and utilization of waste heat. A number of studies have suggested that building-integrated agriculture has the potential to improve overall energy performance of the system (Specht et al 2013). Decentralized residential heating systems in single-family homes make utilization challenging, but specialized building-integrated systems, like the example developed by Seawater Greenhouses, could be a model for smaller-scale units that utilize waste heat and CO₂ on site (Delor 2011). Nevertheless, the model presented by Cerón-Palma et al (2012) of a rooftop greenhouse in Barcelona highlights the challenges of building-integrated UA, as greenhouse heating requirements were not temporally aligned with the times of excess heat within the building; instead, this type of production system may be better suited to colder climates where exhaust CO₂ and heat from boilers/furnaces are more available during winter months. This highlights the need for additional research on how to overcome these types of management issues to support greater resource efficiency.

Historically, UA was a natural part of urban development and, eventually, an essential component of the plans of early urban planning practitioners (Vitiello and Brinkley 2013). However, UA was not a primary objective for planning developed-world public spaces in industrialized food system of 20th century cities. Calls to reconsider the value of UA have been made for decades (e.g. in the pattern language proposed by Alexander et al 1977) and planning for UA, as a result, has returned. The success of UA re-adoption in urban design is demonstrated by the Carrot City Initiative (Gorgolewski et al 2017) which facilitates discussions on urban design for food production. These and other resources can help to increase the sophistication of food planning in a more cyclical urban ecosystem.

Planners can open up or create space to enable the upscaling of UA in either building-integrated systems or new/existing green space. For example, parks could be redeveloped from being merely aesthetically-pleasing recreational landscapes to be more functional, with edible productivity through the incorporation of fruit trees and community gardens. Inventories of suitable public and private vacant land could be identified for UA use through geomatic methods (McClintock et al 2013). Municipal support for training in the harvest and processing of crops could increase the public’s awareness of the resources embodied within the food they consume and minimize and potentially minimize crop waste. Processing infrastructure, such as fruit presses or preserving facilities, could be situated within the park’s borders. By-laws could be put in place to incentivize rooftop UA, as has been done with green roofs in some cities (e.g. Toronto and Chicago; Loder 2014)

As mentioned previously, UA expansion could lead to local increases in polluted run-off. This may require the implementation of by-laws restricting fertilizer or pesticide application, storm water remediation/mitigation measures, and out-reach to inform citizens of health and environmental implications of agriculture. As well, inventories of UA and surveys of practices coupled with geographic information systems could help planners identify potential hotspots for runoff, odors, or other impacts.

Human labor is an abundant urban resource that is anticipated to become more available in cities as trends of urbanization and automation progress. Smaller-scale agricultural systems have the potential to utilize this labor, as they tend to be more labor intensive than conventional mechanized open-field agriculture. As well, the integration of UA in buildings and the application of advanced production approaches (i.e. soilless operations) require specialized training during design, construction and operation, creating high-skilled employment opportunities. The impacts on food prices by shifting to small-scale UA systems is unclear; the 2012 US agricultural census suggests that hired and contract farm labor contributed to only 10.2% of total farm production expenses, though it is suggested that this would vary substantially by crop raised and potentially less mechanized/automated systems (US Department of Agriculture 2014, USDA ERS 2014). The recreational utility realized by those pursuing UA as a leisure activity could reduce the net increase in costs (i.e. people providing free labor in pursuit of UA as a hobby); further, multiple non-monetary benefits (civic engagement, social cohesion, food security) have been recognized, enabling a scenario where broad public benefits of UA can be realized, coupled with an understanding of its effects on health and the environment (Chen 2012, Horst et al 2017).

A number of issues may inhibit efforts to scale up UA, including land scarcity (Martellozzo et al 2014), UA’s uncertain contribution to food security (Ward 2015), environmental impacts of decentralized production (Nicholson et al 2015, Coley et al 2009), and management of new sources of food waste (Levis and Barlaz 2011, Forkes 2007, Smil 2004). Avoiding unintended consequences and continued inefficiency in the food system through urban production requires a planning approach that coordinates input streams, reduces potential for waste and enables co-location to mitigate growth in transportation demand. Foley et al (2011) suggest that efforts to meet the food needs of the rising global (urban) population face substantial challenges to environmental protection. Further, resource demands of all urban food consumption far exceeds the resources that can be provided within city boundaries and moving towards this goal could create new local resource stresses; for example, Ramaswami et al (2017) demonstrate this situation for New Delhi’s water demand, where water used for food production represented 72% of urban-related withdrawals (in turn, only 14% of these water withdrawals was provided within the city’s boundary).

We argue that an industrial ecological approach to UA has the potential to slow land use change (through the intensification of production), increase crops yields (by increasing management intensity), increase resource efficiency (through co-location of inputs from waste streams), and encourage low-carbon diets (through increased access to fresh produce; Wakefield et al 2007, Schafft et al 2009). However, proximity alone are not a guarantee for success of eco-industrial UA; Gibbs and Deutz (2007) review a number of unsuccessful industrial ecological case studies and interview participants in these and find that results often do not match objectives. However, with an incremental planning approach, improved networking to develop trust and cooperation, and targeted policy interventions by municipalities could improve the success of industrial ecological approaches.

Implications of UA on production inputs, food waste and transportation (of both labor and food products) are dependent on UA approaches taken. As an illustration, this will be influenced by the production practices of UA practitioners, efficiency of distribution systems, public and active transportation options for accessing UA sites, producer and retail practices for food disposal, and local attitudes towards food waste. All of these require further study within each local context.